The impacts of warming on the toxicity of carbon nanotubes in mussels

Marine Environmental Research 145 (2019) 11–21

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The impacts of warming on the toxicity of carbon nanotubes in mussels a

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Madalena Andrade , Lucia De Marchi , Carlo Pretti , Federica Chiellini , Andrea Morelli , Etelvina Figueiraa, Rui J.M. Rochaa, Amadeu M.V.M. Soaresa, Rosa Freitasa,∗

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Department of Biology & Center for Environmental and Marine Studies (CESAM), University of Aveiro, 3810-193, Aveiro, Portugal Department of Veterinary Sciences, University of Pisa, San Piero a Grado, Pisa, 56122, Italy Department of Chemistry and Industrial Chemistry, University of Pisa, Udr INSTM Pisa, Pisa, 56126, Italy d Consortium for the Interuniversity Center of Marine Biology and Applied Ecology, Livorno, Italy b c

ARTICLE INFO

ABSTRACT

Keywords: Mussels Nanoparticles Oxidative stress Metabolism Temperature increase

With the increased production and research on nanoparticles, the presence of carbon nanotubes (CNTs) in aquatic systems is very likely to increase. Although it has been shown that CNTs may cause toxicity in marine organisms, to our knowledge, the possible impacts under global temperature increase is still unknown. For this reason, biochemical and physiological impacts induced in Mytilus galloprovincialis due to the presence of functionalized multi-walled CNTs (f-MWCNTs) and increased temperature were investigated in the present study. The mussels exposed to increased temperature alone presented higher metabolic capacity and expenditure of glycogen as an energy resource to fuel up defense mechanisms and thus preventing oxidative damage. Contrarily, organisms exposed to f-MWCNTs alone seemed not stressed enough to demonstrate differences in the metabolism capacity. Furthermore, f-MWCNTs seemed not able to significantly activate their antioxidant and biotransformation enzymes, which in turn may led to oxidative damage in the cells especially when organisms were exposed to a warmer temperature. In fact, at higher temperature, the antioxidant response of organisms exposed to f-MWCNTs was not effective and oxidative damage levels were observed. Nevertheless, no additive or synergetic effects were observed when mussels were exposed to both stressors simultaneously.

1. Introduction The production and research of nanoparticles (NPs) has been increasing in the last years mainly due to their characteristics, including dimensions between 1 and 100 nm (EC, 2011) and specific mechanical, optical, electrical and magnetic properties (Dowling et al., 2004). Particularly, carbon nanoparticles (CNPs) have assumed an important role due to their unique physical and chemical characteristics, presenting a diversity of applications, including in medicine, electronics, energy, food and agriculture sectors (Köhler et al., 2008; Muller and Nowack, 2008; Solarskaciuk et al., 2014; Vlasova et al., 2016; Wu et al., 2013). However, its use in daily life products just as in paints, energy storage and wastewater treatments makes possible their release to the environment (Keller et al., 2013; Mitrano et al., 2015; Petersen et al., 2011). Among the most important CNPs are carbon nanotubes (CNTs) (Liné et al., 2017). Because these NPs easily aggregate in solutions, especially in saltwater (Kataoka et al., 2016), they are commonly functionalized through chemical modifications making them more dispersible and increasing the success of their application (Sun et al., 2002). Thus, the widespread use of CNTs may significantly increase their release into the environment (Potočnik, 2011; ∗

Sanchez et al., 2012). Although the environmentally relevant concentrations (ERCs) of CNTs in water are in the μg/L to ng/L range (Sun et al., 2016), the predicted environmental concentrations (PECs) of CNTs in aqueous systems reported from the most recent literature was projected to approximately 0.001–1000 μg/L (Potočnik, 2011; Sanchez et al., 2012; Zhang et al., 2017). It is already known that in the aquatic environment CNTs can be easily accumulated by the aquatic biota through body surface, digestive and respiratory system (Jackson et al., 2013). Recent studies already have demonstrated that higher water dispersibility of functionalized CNTs (as for example with surface areas containing carboxyl groups) induce higher levels of toxicity to biological systems than the nonfunctionalized ones due to the presence of higher amorphous carbon fragments in comparison to the other form of insoluble CNTs (Arndt et al., 2013; De Marchi et al., 2018a, 2018b; Kataoka et al., 2016). It has been suggested that suspension-feeding invertebrates as bivalves may represent a unique target group for nanoparticles (Moore, 2006). In fact, these organisms have developed processes for the cellular internalization on NPs (Canesi et al., 2010b). However, the impacts of different NPs in bivalves were already demonstrated, particularly for CNTs, which by their dissolution and/or structure, induces immunotoxicity, oxidative stress and

Corresponding author. E-mail address: [email protected] (R. Freitas).

https://doi.org/10.1016/j.marenvres.2019.01.013 Received 23 December 2018; Received in revised form 25 January 2019; Accepted 27 January 2019 Available online 29 January 2019 0141-1136/ © 2019 Elsevier Ltd. All rights reserved.

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cellular injuries, membrane and DNA damage (Rocha et al., 2015). Some specific cases observed are for example: i) induction of inflammatory processes, antioxidant enzymes activity and cytotoxicity in Mytilus galloprovincialis mussels (Canesi et al., 2008, 2010a; Moore et al., 2009); ii) impairment of lysosomal stability in circulating blood cells and lipid peroxidation in Mytilus edulis mussels (Kádár et al., 2010), iii) reduction of the survival or growth of the mussel Villosa iris, with presence of CNTs in the gut (Mwangi et al., 2012); iv) histopathological changes in the epithelium and swelling of the connective tissue of intestine, digestive gland, and gills of the mussel Modiolus modiolus (Anisimova et al., 2015); v) and decrease of mussel Mytilus coruscus settlement (> 50%) (Yang et al., 2016). Beside chemical contamination, bivalves are also exposed to the increase in temperature as a result of global warming, observed since the mid-20th century and close related to the cumulative emissions of anthropogenic greenhouse gases (IPCC, 2014). In particular, the continuous increase of atmospheric CO2 until the end of the 21st century is considered one of the most important contributors to global warming, with projected increases ranging between 1.0 and 4.0 °C of the global mean air temperature (IPCC, 2007). This increase of temperature may cause negative effects in aquatic invertebrates, since those are particularly sensitive to thermal stress due to their ectothermic biology (Pörtner, 2010). In fact, different studies already revealed that when temperature exceeds the organism's thermal tolerance range, physiological and molecular perturbations may happen including in individuals' growth and reproduction (Pörtner and Knust, 2007; Boukadida et al., 2016). Furthermore, the decrease of aerobic capacity, metabolic rate and respiratory capacity were also observed in organisms exposed to warming conditions (Jansen et al., 2009; Pörtner et al., 2005; Pörtner, 2010; Velez et al., 2017). Oxidative stress may as well occur, since warming can enhance the production of reactive oxygen species (ROS) in the cells (Kefaloyianni et al., 2005; Verlecar et al., 2007). Such induction of oxidative stress was already demonstrated in clams (R. philippinarum, R. decussatus) and mussels (M. galloprovincialis, M. coruscus) (Attig et al., 2014; Banni et al., 2014; Hu et al., 2015; Nardi et al., 2017; Velez et al., 2017). The mussel species M. galloprovincialis (Lamarck, 1819) presents a wide distribution and abundance, with known sedentary lifestyle, tolerance to a wide range of environmental conditions and capacity to accumulate and reflect toxic impacts from pollutants, making them widely used as a sentinel and bioindicator species (Banni et al., 2014; Catsiki and Florou, 2006; Faggio et al., 2016; Kristan et al., 2014; Oliveira et al., 2017; Viarengo et al., 2007; Wang et al., 1996). Although the increase of temperature is known to cause physiological and biochemical effects in mussels (Anestis et al., 2007; Coppola et al., 2017, 2018; Gestoso et al., 2016; Kamel et al., 2012), and some information is known about the impacts of functionalized CNTs in M. galloprovincialis (Al-Shaeri et al., 2013; Canesi et al., 2008, 2010a; Miller et al., 2015; Moore et al., 2009; Moschino et al., 2014), no information is known about the effects of the combination of both stressors. Within this context, the present study aimed to evaluate the physiological and biochemical alterations imposed by the presence of functionalized multi-walled CNTs (f-MWCNTs) and increased temperature, both acting alone and in combination, evaluating if increased temperature enhances the toxicity of the MWCNTs.

Furthermore, mussels were kept under continuous aeration during a 12 h light: 12 h dark photoperiod, and fed three times (day 3, 5 and 7 of the acclimation period) with 107 cells/animal of AlgaMac Protein Plus (Aquafaune Bio-Marine Inc., Hawthorne, USA). After this period, organisms were distributed into different aquaria (20 L seawater, salinity 35), with a total of 3 aquaria per condition with 6 organisms in each (total of 18 organisms per condition). The treatments tested were: control temperature (18 °C) without f-MWCNTs (Temp CTL); control temperature (18 °C) with f-MWCNTs (Temp CTL + f-MWCNTs); temperature 21 °C without f-MWCNTs (Temp); temperature 21 °C with f-MWCNTs (Temp + f-MWCNTs). Aquaria were placed in two different climatic rooms to enable the temperature levels of 18 ± 1.0 °C (control temperature) and 21 ± 1.0 °C (increased temperature). The water in the aquaria placed at 21 °C room increased gradually the temperature along the first day of exposure. The selection of the CNTs was based on their different physical and chemical properties and different behavior in the water media (aggregation/disaggregation, adsorption/desorption, sedimentation/resuspension and dissolution). Specifically, f-MWCNTs are more stable in salt water media in comparison to pristine MWCNTs as a consequence of the introduction of oxygen-containing groups on the surface (COOHMWCNTs). These groups ionize in water charging the oxygen atoms negatively and in aqueous phase the electrostatic repulsive forces between negative surface charges of the oxygen-containing groups can lead to stability of oxidized CNTs in the water column, increasing the availability of these materials for the organisms (Peng et al., 2009). Moreover, the selection of CNTs was also based on their wide industrial applications (see specifications MWCNTs-COOH: TNMC1 series, http://www. timesnano.com) and, therefore, potential presence in aquatic systems. The concentration of f-MWCNTs used was chosen: i) following previous works carried out by De Marchi et al. (2017b; 2017c; 2018b) in invertebrate species, namely bivalves R. philippinarum and polychaetes D. neapolitana and H. diversicolor, where 0.01 mg/L was the lowest concentration to induce observable physiological changes and ii) in accordance with the PECs of CNTs in aqueous systems (0.001–1000 μg/L). The control temperature of 18 ± 1.0 °C was selected considering the average temperature of the sampling area during September (IPMA, 2017). The temperature of 21 °C was selected to simulate warming conditions, taking in account the annual range of average temperatures (13.4–22.9 °C) for M. galloprovincialis habitats in Ria de Aveiro (Coelho et al., 2014; Santos et al., 2009; Velez et al., 2015) and bearing in mind the predicted increase of temperature from 1.0 °C to 4.0 °C (IPCC, 2007). During the experimental period of 14 days, organisms were fed three times per week with 107 cells/animal of AlgaMac Protein Plus (Aquafaune Bio-Marine Inc., Hawthorne, USA). After 7 days of beginning of the experiment seawater was renewed, re-establishing seawater characteristics, including salinity, temperature, pH and f-MWCNTs concentration. An experimental period of 14 days was selected taking in account previous studies in mussels (Andrade et al., 2018; Hu et al., 2015; Huang et al., 2018; Letendre et al., 2011; Verlecar et al., 2007) which observed physiological and biochemical changes during this period. At the end of the experimental period of 14 days, the organisms were immediately frozen at −80 °C until further analysis with the exception of two organisms per aquarium which were immediately used for respiration rate determination. Water samples were taken right before of seawater renewal to characterize MWCNTs in the water column.

2. Methodology 2.1. Sampling and experimental conditions Mytillus galloprovincialis mussels were collected in September 2017 at the Mira Channel (Ria de Aveiro, a coastal lagoon at the northwest of Portugal) during low tide. After sampling, organisms were transported to the laboratory and placed in aquaria for depuration and acclimation to laboratory conditions for 7 days. Aquaria were filled with artificial seawater (salinity 35 ± 1), made it with deionized water and artificial salt (Tropic Marin®SEA SALT from Tropic Marine Center). During the acclimation period, organisms were maintained at 18 °C ± 1.0 °C and pH 8.0 ± 0.1 to resemble estuarine conditions.

2.2. MWCNTs characterization The functionalized CNTs (f-MWCNTs) were produced via catalytic carbon vapor deposition (CCVD) process. The carbon nanotubes were purchased from Times Nano: Chengdu Organic Chemicals Co. Ltd., Chinese Academy of Sciences (MWCNTs-COOH: TNMC1 series, http:// www.timesnano.com). The manufacturer specifications are: diameter 2–5 nm; length 10–30 μm; carbon purity 98%; surface area 400 m2/g; amorphous carbon 8–10% and -COOH 3.86 wt%. 12

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A stock solution of 50 mg/L f-MWCNTs concentration, using the same artificial seawater arranged for the aquaria, was prepared for the concentration of MWCNTs used in this study (0.01 mg/L). A stirrer was used to facilitate the dispersion of f-MWCNTs in the aquaria before pipetting the solution and the reoxygenation system of the water was used to avoid its complete precipitation. The average size distribution and the polydispersity index (PDI) of f-MWCNT suspensions in seawater in each exposure condition was analyzed by dynamic light scattering (DLS), using a Delsa™ NanoC Particle Size Analyser (Beckman Coulter) for particles characterization. Immediately before water renewal water samples (15 mL) were collected for characterization analysis. Measurements were carried out on 1 mL of suspension and each analysis was repeated three times. Due to the inherent heterogeneity and colloidal instability of the analyzed samples, DLS analyses were repeated several times to ensure reproducible results. Undetected colloidal material at the end of each measurement was designated in the Table 1 as Invalid data (I.d.) while “no data” represented invalid data (I.d.) results in 3 out of 5 samples. Intensity distributions were obtained by analyzing the autocorrelation functions through the Contin algorithm which is particularly ideal for polydisperse and multimodal systems. The cumulant method was used to determine the hydrodynamic radius and polydispersity index (PDI) of the analyzed dispersions on three replicates (1A,1B and 1C) of each sample collected after a week (t7) of the experimental period. Moreover, a control water samples without animals (t0) was used observing the natural behavior of f-MWCNTs in the salt water media.

to determinate mussels’ lipid content (LIP). The dry tissue from the whole organism was homogenized with a mortar and a pestle, being divided in 15 mg aliquots and stored for LIP quantification. For all other parameters except LIP, shells of the frozen organisms (four per aquarium, twelve per condition) were removed and the frozen whole soft tissue was pulverized using a mortar and pestle with liquid nitrogen. The homogenized tissue of each organism was then distributed in 0.5 g aliquots for further quantifications. For each biochemical parameter, the extraction of the supernatant was performed with a specific buffer using a proportion of 1:2 (w/v) (see Andrade et al., 2018; De Marchi et al., 2017b; Mesquita et al., 2014). Tissue samples were homogenized during 1 min using a TissueLyser II (Qiagen), being after centrifuged for 20 min at 4 °C and 10000 g or 3000 g depending of the biomarker. The supernatants were then stored at −80 °C or immediately used to determine: electron transport system (ETS) activity; glycogen (GLY) content; lipid peroxidation (LPO) and protein carbonylation (PC) levels; activity of antioxidant enzymes (superoxide dismutase, SOD; glutathione peroxidase, GPx and catalase, CAT) and glutathione-S-transferases (GSTs). Two replicates per sample were used for the assessment of each biochemical parameter. 2.4.1. Metabolic capacity The ETS activity was determined based on the method of King and Packard (1975) and modifications by Coen and Janssen (1997). Absorbance was measured during 10 min at 490 nm with intervals of 25 s and the extinction coefficient (ε) of 15,900 M−1cm−1 was used to calculate the amount of formazan formed. Results were expressed in nmol min per g of fresh weight (FW).

2.3. Biological responses: physiological parameters 2.3.1. Condition index The Condition Index (CI) was calculated taking in account that this parameter can give an indication of the general physiological status of the animals (Andral et al., 2004). After the experimental period (14 days), the soft tissues of six frozen organisms per condition (two per aquarium, six per condition) were carefully separated from the shells. Both separated shells and tissues were placed in an oven at 60 °C for 48 h. After this period, both dry soft tissues and shells were weighed and CI calculated, corresponding to the ratio between the dry weight of softs tissues and the dry weight of shell x 100 following Matozzo et al. (2012). The dry tissue was stored and used for lipid quantification.

2.4.2. Energy reserves The GLY content was quantified following the sulfuric acid method (Dubois et al., 1956), using 8 glucose standards in the concentration range of 0–10 mg/mL in order to obtain a calibration curve. Absorbance measurement was read at 492 nm after being incubated for 30 min at room temperature. Results were expressed in mg per g of FW. The LIP content was determined based in the methods developed by Folch et al. (1957) and Cheng et al. (2011). A standard curve was estimated using cholesterol standards (0–100%). The absorbance was read at 520 nm, after 1 h of color development in the dark at the room temperature. Results were expressed in percentage per mg DW.

2.3.2. Respiration rate After 14 days of exposure respiration rate (RR) was measured in six mussels per condition (two per aquarium). Measurements were performed by simple static respirometry, in two organisms of the same aquarium per respirometric chamber filled to its maximum capacity (1 L) to avoid the formation of air bubbles, with the same seawater used during the experimental period. For 2 h, organisms were placed in the dark and oxygen concentrations were measured every 15 min with a multi-channel fiber optic oxygen meter (Multi channel oxygen meter PreSens) for simultaneous read-out, equipped with an oxygen sensor spot glued to its wall using silicon paste. Measurements were recorded using the software PreSens Measurement Studio 2. Animals were allowed to acclimate for 30 min to avoid the influence of manipulation on RR. After this time, and in fully oxygenated medium, chambers were air-tight sealed and RR was recorded as a function of declining O2 concentration (mg/L) over time. Twenty-two measurements, corresponding to twenty-two chambers with two individuals each, were carried out at the same time (including a blank) and data were recorded in 3 s intervals between chambers. Organisms were posteriorly dried and weighed. Respiration rate was expressed in mg O2 consumed per h per g dry weight (DW) and correction to O2 variation in blanks (chambers with no organisms) was employed.

2.4.3. Antioxidant enzymes The activity of SOD was quantified following the method of Beauchamp and Fridovich (1971). The standard curve was formed using SOD standards (0.25–60 U/mL). Samples’ absorbance was read at 560 nm after 20 min of incubation at room temperature. Results were expressed in U per g FW where one unit (U) of enzyme activity corresponds to a reduction of 50% of nitroblue tetrazolium (NBT). The activity of GPx was determined based the method of Paglia and Valentine (1967). Absorbance measurement was read at 340 nm (ε = 6.22 mM−1cm−1) during 5 min in 10 s intervals. Results were expressed in U per g FW where one unit (U) corresponds to the quantity of enzyme which catalyzes the conversion of 1 μmol nicotinamide adenine dinucleotide phosphate (NADPH) per min. The activity of CAT was determined following Johansson and Borg (1988). The standard curve was determined by using formaldehyde standards (0–150 μM). Absorbance was read at 540 nm. Results were expressed in U per g FW where one unit (U) is defined as the formation of 1 nmol formaldehyde per min in this case. The activity of GSTs was quantified following the method described in Habig et al. (1974) protocol with some adaptations performed by Carregosa et al. (2014c). GSTs activity was read at 340 nm (ε = 9.6 mM−1 cm−1). Results were expressed in U per g of FW where U is defined as the amount of enzyme that catalysis the formation of 1 μmol of dinitrophenyl thioether per min.

2.4. Biological responses: biochemical parameters After the 14 days of the experimental period, the dry tissue of six organisms per condition (used previously for the CI and RR) were used 13

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2.4.4. Oxidative damage Levels of LPO were measured according with the method described in Ohkawa et al. (1979) and with modifications stated by Carregosa et al. (2014b). Absorbance was read at 535 nm (ε = 156 mM−1 cm−1) and results expressed in nmol of MDA equivalents formed per g FW. Levels of PC were quantified following the DNPH alkaline method described by Mesquita et al. (2014). Absorbance was read at 450 nm (ε = 22.00 M−1 cm−1). Results being expressed in nmol of protein carbonyls groups formed per g FW.

Table 1 Dynamic light scattering (DLS) data of size (nm) and polydispersity index (PDI) of water samples exposed to control temperature (18 °C) with functionalized multi-walled carbon nanotubes (f-MWCNTs) (CTL + f-MWCNTs) and increase temperature (21 °C) with f-MWCNTs (Temp + f-MWCNTs) at the time 0 (control water samples without animals) and after one week of exposure (t7). Samples 1A, 1B, 1C: experiment replicates carried out at the cited conditions contaminated with 0.01 mg/L of f-MWCNTs. I.d.: “Invalid data” (not detected colloidal material into the analyzed sample at the end of 120 acquisitions); n.d.: “no data” (invalid data (I.d.) results in 3 out of 5 samples).

2.5. Data analysis Due to a lack of homogeneity of variance, CI, RR, ETS, GLY, LIP, SOD, GPx, CAT, GSTs LPO and PC were separately submitted to a non-parametric permutational analysis of variance (PERMANOVA Add-on in Primer v7) with a two factors design: temperature condition (exposed to control temperature (TempCTL) and exposed to increased temperature (Temp)) as factor 1 and contaminated condition as factor 2. PERMANOVA main test was performed to test the effect of temperature condition, contaminated condition and the interaction between these two factors on each biomarker. PERMANOVA main tests were considered significant when p ≤ 0.05 and followed by PERMANOVA pair-wise tests. Pair-wise tests were used to test the effect of temperature condition within each contaminated condition and the effect of contaminated condition within each temperature condition, testing the hypotheses: i) No significant differences exist among temperature within noncontaminated contaminated or within contaminated organisms; ii) No significant differences exist between non-contaminated and contaminated for each temperature. PERMANOVA pair-wise tests results are represented in figures with lower case letters and in the main text by p-values.

Samples

Size (nm)

PDI

CTL + f-MWCNTs t0 CTL + f-MWCNTs t7_1A CTL + f-MWCNTs t7_1B CTL + f-MWCNTs t7_1C Temp + f-MWCNTs t0 Temp + f-MWCNTs t7_1A Temp + f-MWCNTs t7_1B Temp + f-MWCNTs t7_1C

1109.5 n.d. n.d. n.d. n.d. 1123.4 2716.2 1936.0

0.794 – – – – 0.446 1.150 0.816

observed between both temperature treatments (Fig. 1B). Comparing contaminated and non-contaminated organisms, for the control temperature no significant differences were observed (Fig. 1B). For increased temperature, although with no significant differences, lower RR values were observed in contaminated mussels in comparison non-contaminated mussels (Fig. 1B). 3.3. Biochemical parameters 3.3.1. Metabolic capacity Concerning electron transport system (ETS) activity and comparing both temperature treatments, significantly higher values were observed in organisms exposed to warming conditions, both for contaminated and non-contaminated mussels (Fig. 2A). No significant differences were observed between non-contaminated and contaminated organisms for each tested temperature (Fig. 2A).

3. Results 3.1. MWCNTs characterization As reported in the literature, Dynamic light scattering (DLS) was used to measure the mean size (nm) and the polydispersity index (PDI) of simulated seawater samples collected from aquaria containing mussels (Mytilus galloprovincialis). Here DLS analyses were repeated several times to ensure reproducible results due to the inherent heterogeneity and colloidal instability of the analyzed samples. All the results obtained from the three replicates (1A,1B and 1C) after one week of exposure (t7) didn't evidence the presence of dispersed materials in seawater samples collected from aquaria where organisms were subjected to control temperature (18 °C) and f-MWCNTs while samples subjected to increased temperature (21 °C) were contaminated by micro-sized suspensions (maximum size diameter 2716.2 nm) (Table 1).

3.3.2. Energy reserves The glycogen (GLY) content was significantly lower in mussels exposed to increased temperature in comparison to control temperature, both for non-contaminated and contaminated organisms (Fig. 2B). Comparing noncontaminated and contaminated organisms, no significant differences were observed regardless the temperature tested (Fig. 2B). Lipid (LIP) content showed no significant differences between mussels exposed to control and increased temperature, both for noncontaminated and contaminated mussels (Fig. 2C). No significant differences were observed between contaminated and non-contaminated organisms, for both temperatures (Fig. 2C).

3.2. Physiological parameters

3.3.3. Antioxidant enzymes Concerning superoxide dismutase (SOD) activity and comparing both temperatures no significant differences were observed both for non-contaminated and contaminated mussels (Fig. 3A). Comparing contaminated and non-contaminated organisms, for both temperatures significantly lower SOD values were observed in contaminated mussels (Fig. 3A). The glutathione peroxidase (GPx) activity in non-contaminated organisms was significantly higher at increased temperature in comparison to control temperature, while no significant differences were observed between both temperatures for contaminated mussels (Fig. 3B). Comparing non-contaminated and contaminated organisms, at control temperature no significant differences were observed while at warming conditions non-contaminated mussels showed significantly higher values than the contaminated ones (Fig. 3B). The activity of catalase (CAT) showed no significant differences between non-contaminated and contaminated organisms regardless the tested temperature (Fig. 3C). Comparing non-contaminated and contaminated organisms, no significant differences were observed for

3.2.1. Mortality After the experimental period of 14 days, no mortality was recorded. 3.2.2. Condition index Concerning condition index (CI) values and comparing both temperatures treatments (control temperature-Temp CTL and increased temperature-Temp), for non-contaminated and contaminated organisms no significant differences were observed between temperatures (Fig. 1A). Comparing contaminated and non-contaminated organisms, for each temperature treatment, no statistically differences were observed (Fig. 1A). 3.2.3. Respiration rate The respiration rate (RR) in non-contaminated organisms, although with no significant differences, was higher in mussels exposed to increased temperature in comparison to organisms exposed to control temperature (Fig. 1B). For contaminated organisms no significant differences were 14

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Fig. 1. A: Condition Index (CI) and B: Respiration Rate (RR), in Mytilus galloprovincialis exposed to different conditions (Temp CTL, Temp CTL + fMWCNTs; Temp; Temp + f-MWCNTs). Results are the means + standard errors. White bars represent organisms exposed to control temperature (18 °C) while gray bars represent organisms under warming conditions (21 °C). Different letters represent significant differences (p ≤ 0.05) among conditions.

organisms under both control and increased temperatures (Fig. 3C). Relatively to glutathione s-transferases (GSTs) activity for non-contaminated mussels significantly higher values were observed at control temperature in comparison to organisms under warming condition, while no significant differences were observed between temperature treatments for contaminated mussels (Fig. 4). Comparing contaminated and non-contaminated organisms, for the control temperature significant lower values were observed in contaminated mussels, while no significant differences were observed at increased temperature (Fig. 4).

Fig. 2. A: Electron transport system (ETS) activity; B: Glycogen (GLY) content and C: Lipids (LIP) content, in Mytilus galloprovincialis exposed to different conditions (Temp CTL, Temp CTL + f-MWCNTs; Temp; Temp + f-MWCNTs). Results are the means + standard errors. White bars represent organisms exposed to control temperature (18 °C) while gray bars represent organisms under warming conditions (21 °C). Different letters represent significant differences (p ≤ 0.05) among conditions.

3.3.4. Oxidative damage Lipid peroxidation (LPO) levels in non-contaminated organisms showed significantly lower values at increased temperature compared to control temperature (Fig. 5A). For contaminated organisms no significant differences were observed between both temperature treatments (Fig. 5A). Comparing contaminated and non-contaminated organisms, significantly higher LPO values were observed in contaminated mussels for both temperature treatments (Fig. 5A). Protein carbonylation (PC) showed lower PC values in organisms exposed to increased temperature in comparison to mussels under control temperature, with significant differences only for non-contaminated mussels (Fig. 5B). Comparing contaminated and non-contaminated organisms, significant differences were only observed for mussels under warming conditions (Fig. 5B).

4. Discussion The present study evaluated the physiological and biochemical performance of M. galloprovincialis specimens when exposed to increased temperature and f-MWCNTs exposure, acting alone and in combination, aiming to understand if the physiological and biochemical effects due to warming conditions would be enhanced in organisms 15

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Fig. 4. Glutathione S-transferases (GSTs) activity in Mytilus galloprovincialis exposed to different conditions (Temp CTL, Temp CTL + f-MWCNTs; Temp; Temp + f-MWCNTs). Results are the means + standard errors. White bars represent organisms exposed to control temperature (18 °C) while gray bars represent organisms under warming conditions (21 °C). Different letters represent significant differences (p ≤ 0.05) among conditions.

Fig. 3. A: Superoxide dismutase (SOD) activity; B: Glutathione peroxidase (GPx) activity and C: Catalase (CAT) activity, in Mytilus galloprovincialis exposed to different conditions (Temp CTL, Temp CTL + f-MWCNTs; Temp; Temp + f-MWCNTs). Results are the means + standard errors. White bars represent organisms exposed to control temperature (18 °C) while gray bars represent organisms under warming conditions (21 °C). Different letters represent significant differences (p ≤ 0.05) among conditions.

Fig. 5. A: Lipid peroxidation (LPO) levels and B: Protein carbonylation (PC) levels, in Mytilus galloprovincialis exposed to different conditions (Temp CTL, Temp CTL + f-MWCNTs; Temp; Temp + f-MWCNTs). Results are the means + standard errors. White bars represent organisms exposed to control temperature (18 °C) while gray bars represent organisms under warming conditions (21 °C). Different letters represent significant differences (p ≤ 0.05) among conditions.

et al., 2012; Matozzo et al., 2012; Azpeitia et al., 2016). One of the physiological parameters that may indicate the organisms' general health status is the condition index (CI), which may provide information on bivalves’ growth over a certain period of time (Lucas and Beninger, 1985; Orban et al., 2002). Our findings demonstrated that temperature did not influence significantly mussels CI, although a slight increase was observed at higher temperatures which may result from

exposed to an emergent contaminant. 4.1. Physiological responses The physiology of bivalves has been showing to change not only due to reproductive cycles, but also due to different environmental factors such as increased temperature and the presence of contaminants (Çelic 16

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higher filtration rate and, consequently, higher ingestion of food under this condition. Nevertheless, previous studies developed by Kamel et al. (2014) reported that Mytilus galloprovincialis increased their CI especially under August, corresponding to the warmer month analyzed, suggesting that CI may be affected by water temperature. The present results further showed that CI was not significantly affected by fMWCNTs, regardless the temperature, although a slight decrease was observed in contaminated mussels. Such results may indicate that the low concentration of f-MWCNTs tested could be not enough to induce alterations at this level. Another physiological parameter that can be used to assess alterations induced in organisms as a result of different stressors is the respiration rate (RR) (Gestoso et al., 2016; Freitas et al., 2017; Wang et al., 2015). Our results indicate that under increased temperature organisms tended to increase their respiratory capacity, especially in the absence of f-MWCNTs. Different studies have demonstrated that RR may be dependent on the temperature in bivalves (Bayne and Newell, 1983; Griffiths and Griffiths, 1987). Perna perna mussels showed an increase of RR with the increase of different temperature (15, 20, 25 and 30 °C) both after short and long-term exposures (Resgalla Jr. et al., 2007). Also M. galloprovincialis had their RR enhanced with increasing temperature of 21 °C (Gestoso et al., 2016). Likewise, for the same species Jansen et al. (2009) demonstrated an overall increase of RR with the increase of temperature (17, 24, 27, 31 and 34 °C). In the present study, the increased RR observed with the increase of temperature could be related to the increase of metabolism associated to the enhance of antioxidant defenses to avoid cellular damage under warming stress. Our study also revealed that organisms exposed to CNTs were able to maintain (at control temperature) or even decrease (at increased temperature) their RR, probably by maintaining their valves closed to avoid the accumulation of contaminant and prevent oxidative stress. However, the increase of RR has been demonstrated as a physiological adaptation-response to contaminants exposure (Relexans et al., 1988), namely in the mussel Perna perna from contaminated areas (Resgalla et al., 2010) and in cockles from an Hg contaminated area (Nilin et al., 2012). Nonetheless, De Marchi et al. (2017b) observed that when exposed to 0.01 mg/L of MWCNTs, polychaetes Diopatra neopolitana and Hediste diversicolor did not show any changes in RR in comparison to un-contaminated organisms, but at higher concentration (1.00 mg/L of MWCNTs) an increase of RR in H. diversicolor was observed. Therefore, the fact that in the present study the RR did not varied significantly with contamination may result from the low concentration of f-MWCNTs tested and the short exposure period used, which may not be enough to induce any changes on the RR of mussels.

increase in energy costs and consequently maintenance of the oxidative status. Although the increase of ETS activity in mussels under warming conditions was not augmented in the presence of f-MWCNTs, the toxicity at this condition may also result from the fact that salts present in the water may cause the CNTs to precipitate (Peng et al., 2009), especially during a warmer temperature where water may evaporate and salinity increase. Nevertheless, at each of the tested temperatures our results further revealed that mussels exposed to f-MWCNTs were able to maintain their metabolic capacity compared to non-contaminated mussels. The similar values in ETS activity between the contaminated and noncontaminated organisms may indicate an independency in the respiratory capacity with the presence or absence of MWCNTs, further indicating that the concentration used may have not been stressful enough to alter mussels’ metabolic activity. In fact, at the same concentration level of CNTs (0.01 mg/L), De Marchi et al. (2017c, 2018b) did not recognize any significant difference on the ETS activity in clams Ruditapes philippinarum comparing to un-contaminated clams. Besides energy consumption, the availability of energy reserves such as glycogen (GLY) and lipids (LIP) can be affected by general physiological stressors (Scott-Fordsmand and Weeks, 2000). Our results demonstrated an overall decrease of GLY at increased temperature and the maintenance of the LIP levels regardless the tested condition, revealing that associated with higher ETS activity at higher temperature, mussels used their GLY content probably as a resource of energy to fuel up defense mechanisms while LIP was not the main source of energy for mussels to maintain their health status. Previous studies by Clements et al. (2018) also reported that under increased temperatures M. edulis decreased GLY content while maintaining LIP content. The expenditure of energy reserves in bivalves has been observed as well under another stressful conditions. For instance, clams R. philippinarum were found to mobilize GLY content as a source of energy under salinity stressful conditions (Velez et al., 2016). Similarly, oysters Cassostrea virginica exhibited a partial depletion of GLY and LIP reserves when subjected to stressful salinity and hypercapnic conditions (Dickinson et al., 2012). The present study further demonstrated that exposure to f-MWCNTs did not alter mussels' energy reserves content. Such results may indicate that the exposure to f-MWCNTs was not stressful enough to cause an extra expenditure of energy reserves which could once again be related to the low concentration used (0.01 mg/L) and short experimental period tested. In fact, De Marchi et al. (2018b) observed similar results in R. philippinarum exposed to the same MWCNTs concentration. However, the same authors observed a decrease of GLY content at higher MWCNTs concentrations (0.10 and 1.00 mg/L). Different studies have in fact demonstrated the expenditure of energy reserves in invertebrates exposed to higher concentrations of CNPs (De Marchi et al., 2017a, 2017b, 2017c), suggesting again that the GLY contents’ preservation observed in the present study may result from the low f-MWCNT concentration used which did not result in the expenditure of this reserve.

4.2. Biochemical responses 4.2.1. Metabolic capacity and energy reserves The energy consumption at a mitochondrial level, which can be estimated by the determination of electron transport system (ETS) activity, can give an indication of the organisms’ metabolic status (Coen and Janssen, 1997; Berridge et al., 2005; Fanslow et al., 2001; García-Martín et al., 2014). In the present study, both non-contaminated and contaminated organisms exposed to increased temperature were able to increase their ETS activity, which accompanied higher RR. In mussels the increase of ETS activity with the increase of temperature has been observed in different studies. For example, Doucet-Beaupré et al. (2010) demonstrated an increase of ETS activity in the mussel M. edulis with the increase of temperature. Likewise, Fanslow et al. (2001) observed an increase of ETS activity in the mussel Dreissena polymorpha exposed to warming conditions. Organisms try to adapt to increasing temperature by adjusting the density and functional properties of the mitochondria, which consequently affects the production of ROS and antioxidant defences (Martínez-Álvarez et al., 2005). Therefore, the behaviour observed might be related with the activation of the respiratory chain as an adaptation to a stressful warmer condition being associated with an

4.3. Oxidative stress It is well known that under stressful conditions marine organisms increase the production of reactive oxygen species (ROS) and to eliminate the excess of ROS when may increase their antioxidant defenses, as the activity of superoxide dismutase (SOD) and catalase (CAT) enzymes (Freitas et al., 2016b; Regoli and Giuliani, 2014; Velez et al., 2016). Another mechanism which neutralizes ROS directly is the antioxidant enzyme glutathione peroxidase (GPx) which transforms reduced glutathione (GSH) to oxidized glutathione (GSSG )reducing lipid hydroperoxides (Regoli and Giuliani, 2014). Overall, our findings showed that under increased temperature, non-contaminated mussels slightly increased their antioxidant defenses which may in part explain the decrease observed on lipid peroxidation (LPO) levels; while in the presence of f-MWCNTs organisms were not able to increase their antioxidant defenses probably because the stress caused by the presence of the NPs was not enough to increase the activation of the antioxidant 17

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enzymes or because such enzymes could be at a certain limit inhibited by the presence of f-MWCNTs. Increased SOD activity has been observed in M. galloprovincialis exposed to an increased temperature (21 °C) after 28 days of exposure (Coppola et al., 2017). In the same organism, Kamel et al. (2012) demonstrated that both CAT and LPO increased under 22 °C. In the mussel M. coruscus, except for GPx, an increase of SOD and CAT activities in the mussel under warmer conditions (Hu et al., 2015) was observed. Our results further revealed that organisms exposed to f-MWCNTs did not change the enzymes activity compared to non-contaminated ones, which again may indicate that the low concentration of f-MWCNTs tested in this work was not sufficient to induce alterations in the antioxidant enzymes activity. Also, these results could explain higher LPO and protein carbonylation (PC) levels in contaminated mussels compared to non-contaminated ones, especially under higher temperature. Such results may indicate that f-MWCNTs could inhibit the activity of antioxidant enzymes, which could be reflected in a significant decrease on enzymes activity if the tested concentration of f-MWCNTs was higher. At the same concentration of MWCNTs, De Marchi et al. (2017a) demonstrated that for the polychaetes D. neopolitana and H. diversicolor, CAT and SOD activities did not change. Similarly, De Marchi et al. (2018b) observed in R. philippinarum exposed to 0.01 mg/L of MWCNTs no activation of the SOD, CAT and GPx antioxidant activity. Glutathione-S-transferases (GSTs) are a group of enzymes with a detoxification capacity, catalyzing the conjugation of GSH to xenobiotic substrates to be later expelled by the organism (Regoli and Giuliani, 2014). Our findings indicated that control organisms decreased GSTs activities when exposed to increased temperature, which may indicate that warmer conditions can inhibit GSTs activity. Kamel et al. (2014) observed that GSTs inhibition on M. galloprovincialis was higher in June and August corresponding to the warmer months of the experiment. Balbi et al. (2017) observed in the same organisms, seasonal variations of GSTs activity having lower values during the warmer seasons. The present findings also showed that the decrease of GSTs in contaminated mussels was similar at control and increased temperature. These results showed f-MWCNTs will inhibit the activity of GSTs but above certain limits of stress these enzymes are no longer inhibited. For this reason, when exposed to warming conditions and f-MWCNTS mussels showed similar GSTs levels than contaminated organisms under control temperature. Likewise, clams R. philippinarum exposed to f-MWCNTS (0.10 mg/L) showed to inhibit GSTs activity (De Marchi et al., 2018b). Furthermore, Anisimova et al. (2015) demonstrated a decrease of GSTs activity in Crenomytilus grayanus mussels exposed to MWCNTs (100 mg/L). It has been described that the over production of ROS in marine bivalves can cause oxidative damage in the membrane lipids if antioxidant defenses are not efficient in ROS elimination (Carregosa et al., 2014a; Freitas et al., 2016a; Liu et al., 2007; Lushchak, 2011; Matozzo et al., 2012; Silva et al., 2005). The present results showed that when exposed to increased temperature LPO levels decreased in non-contaminated mussels in comparison to organisms from the control temperature, while in contaminated mussels LPO was similar at both tested temperatures. These findings may indicate that under warming conditions non-contaminated organisms were able to avoid cellular damages by increasing the antioxidant defenses. However, these protective mechanisms were not so effective in organisms exposed to f-MWCNTs, which may indicate a higher stressful condition in these organisms and, thus, higher ROS production occurred with associated cellular damage. Kamel et al. (2012) demonstrated that when M. galloprovincialis were exposed for 7 days to 20 °C no differences were observed in LPO levels, while increasing significantly at 22 °C. The present findings further demonstrated that at both temperatures contaminated organisms increased their LPO levels in comparison to non-contaminated mussels revealing high stress levels and increased production of ROS in the presence of f-MWCNTs. The increase of LPO levels in contaminated organisms was more pronounced under warming conditions. The presence of CNTs at the warmer temperature, as revealed by DLS analysis and explained by a probable precipitation

and aggregation in salt water may have resulted in the uptake of this NPs by the organisms with intracellular accumulation which possibly enhanced the oxidative degradation of lipids. In fact, the aggregation of NPs may alter their biological effects by affecting ion release from the surface, their reactive surface area and the consequence mode of cellular uptake of NMs together bringing subsequent biological responses in the organisms (Hotze et al., 2010). Moreover, a study conducted by Ward et al. (2009) showed that generally mussels and oysters capture and ingest more efficiently NPs that are incorporated into aggregates compared to those freely suspended in the water column due to the fact that aggregates are broken down by the action of cilia on the gills and labial palps, being the constituent particles ingested. Previous studies conducted by Anisimova et al. (2015) demonstrated an increase of LPO levels in C. grayanus exposed to 12–14 nm diameter MWCNTS (100 mg/L) for 48 h. Additionally, De Marchi et al. (2017b, 2017c, 2018b) showed an increase of LPO in clams for the same MWCNTs concentration (0.01 mg/ L) in R. phillipinarum and polychaetes D. neopolitana and H. diversicolor after a 28 days exposure period. ROS may alter essential cellular functions though reversible or irreversible post-translational modifications (PTM) which is the last case may inactivate critical proteins (Sultan et al., 2018). PC is one of the most common type of PTM triggered by oxidative stress, which is a process promoted by ROS through the oxidation of proteins (Cattaruzza and Hecker, 2008; Suzuki et al., 2010). Our results revealed that under increased temperature, PC levels either slightly decreased in non-contaminated mussels or were maintained in contaminated organisms. These results indicate that mussels exposed to increased temperature seemed to be able to avoid the carbonylation of proteins by activating the antioxidant enzymes due to a higher ROS production under this condition. In the crabs Scylla serrata it was observed an increase of carbonyls in the muscles of mud crabs collected in summer in comparison to organisms captured during winter (Paital and Chainy, 2013). The present findings also highlighted that no alteration on PC level in mussels were observed in contaminated organisms compared to non-contaminated organisms at the control temperature. Nonetheless, at increased temperature, contaminated organisms showed higher PC levels when compared to noncontaminated organisms, suggesting that antioxidant levels were not effective in protecting organisms against the over production of ROS. Marisa et al. (2016) showed no significant differences of protein carbonyl content neither LPO levels in R. philippinarum exposed during 7days to zinc oxide nanoparticles (1 μg/L nZnO, 10 μg/L nZnO, 10 μg/L ZnCl2), suggesting that antioxidant defenses were enough to cope with the increase of oxidative damage and protect the cells. Overall, the present study demonstrated that exposure to increased temperature or f-MWCNTs may affect the physiological and biochemical performance of M. galloprovincialis, but no additive effects were observed in mussels under both stressors. Acknowledgments Lucia De Marchi benefited from PhD grant (SFRH/BD/101273/ 2014) given by the National Funds through the Portuguese Scleroderma Foundation, supported by FSE and Programa Operacional Capital Humano (POCH) e da União Europeia. Rui J. M. Rocha was supported by a Postdoc scholarship (SFRH/BPD/99819/2014). Rosa Freitas benefited from a Research position funded by Integrated Programme of SR&TD “Smart Valorization of Endogenous Marine Biological Resources Under a Changing Climate” (reference Centro-01-0145-FEDER000018), co-funded by Centro 2020 program, Portugal 2020, European Union, through the European Regional Development Fund. Thanks are also due, for the financial support to CESAM (UID/AMB/50017), to FCT/MEC through national funds, and the co-funding by the FEDER, within the PT2020 Partnership Agreement and Compete 2020; Programa Operacional Competitividade e Internacionalização FEDER (POCI-01-0145-FEDER-028425) BISPECIAl - BIvalves under Polluted Environment and ClImate change. 18

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References

condition index, and biochemical composition of mussels (Mytilus galloprovincialis Lamarck, 1819) in Sinop, South of the Black Sea. J. Aquat. Food Prod. Technol. 21, 198–205. https://doi.org/10.1080/10498850.2011.589099. Cheng, Y., Zheng, Y., VanderGheynst, J.S., 2011. Rapid quantitative analysis of lipids using a colorimetric method in a microplate format. Lipids 46, 95–103. https://doi. org/10.1007/s11745-010-3494-0. Clements, J.C., Hicks, C., Tremblay, R., Comeau, L.A., 2018. Elevated seawater temperature, not pCO2, negatively affects post-spawning adult mussels (Mytilus edulis) under food limitation. Conserv. Physiol. 6 (1). https://doi.org/10.1093/conphys/ cox078. Coelho, J.P., Pato, P., Henriques, B., et al., 2014. Long-term monitoring of a mercury contaminated estuary (Ria de Aveiro, Portugal): the effect of weather events and management in mercury transport. Hydrol. Process. 28, 352–360. https://doi.org/10. 1002/hyp.9585. Coen, W.M.D., Janssen, C.R., 1997. The use of biomarkers in Daphnia magna toxicity testing. IV. Cellular Energy Allocation: a new methodology to assess the energy budget of toxicant-stressed Daphnia populations. J. Aquatic Ecosyst. Stress Recovery 6, 43–55. http://doi.org/10.1023/A:1008228517955. Coppola, F., Almeida, Â., Henriques, B., Soares, A.M.V.M., Figueira, E., Pereira, E., Freitas, R., 2017. Biochemical impacts of Hg in Mytilus galloprovincialis under present and predicted warming scenarios. Sci. Total Environ. 601–602, 1129–1138. https:// doi.org/10.1016/j.scitotenv.2017.05.201. Coppola, F., Almeida, Â., Henriques, B., Soares, A.M.V.M., Figueira, E., Pereira, E., Freitas, R., 2018. Biochemical responses and accumulation patterns of Mytilus galloprovincialis exposed to thermal stress and Arsenic contamination. Ecotoxicol. Environ. Saf. 147, 954–962. https://doi.org/10.1016/j.scitotenv.2017.05.201. De Marchi, L., Neto, V., Pretti, C., Figueira, E., Chiellini, F., Morelli, A., Soares, A.M.V.M., Freitas, R., 2017a. The impacts seawater acidification on Ruditapes philippinarum sensitivity to carbon nanoparticles exposure. Environ. Sci. Nano 4, 1692170. https:// doi.org/10.1039/c7en00335h. De Marchi, L., Neto, V., Pretti, C., Figueira, E., Chiellini, F., Soares, A.M.V.M., Freitas, R., 2017b. Physiological and biochemical responses of two keystone polychaete species: Diopatra neapolitana and Hediste diversicolor to Multi-walled carbon nanotubes. Environ. Res. 154, 126–138. https://doi.org/10.1016/j.envres.2016.12.018. De Marchi, L., Neto, V., Pretti, C., Figueira, E., Chiellini, F., Soares, A.M.V.M., Freitas, R., 2017c. The impacts of emergent pollutants on Ruditapes philippinarum: biochemical responses to carbon nanoparticles exposure. Aquat. Toxicol. 187, 38–47. https://doi. org/10.1016/j.aquatox.2017.03.010. De Marchi, L., Neto, V., Pretti, C., Figueira, E., Chiellini, F., Morelli, A., Soares, A.M.V.M., Freitas, R., 2018a. Effects of multi-walled carbon nanotube materials on Ruditapes philippinarum under climate change: The case of salinity shifts. Aquat. Toxicol. 199, 199–211. https://doi.org/10.1016/j.aquatox.2018.04.001. De Marchi, L., Neto, V., Pretti, C., Figueira, E., Chiellini, F., Morelli, A., Soares, A.M.V.M., Freitas, R., 2018b. Toxic effects of multi-walled carbon nanotubes on bivalves: Comparison between functionalized and nonfunctionalized nanoparticles. Sci. Total Environ. 622–623, 1532–1542. https://doi.org/10.1016/j.scitotenv.2017.10.031. Dickinson, G.H., Ivanina, A.V., Matoo, O.B., Pörtner, H.O., Lannig, G., Bock, C., Beniash, E., Sokolova, I.M., 2012. Interactive effects of salinity and elevated CO2 levels on juvenile eastern oysters, Crassostrea virginica. J. Exp. Biol. 215, 29–43. https://doi. org/10.1242/jeb.061481. Doucet-Beaupré, H., Breton, S., Chapman, E.G., Blier, P.U., Bogan, A.E., Stewart, D.T., Hoeh, W.R., 2010. Mitochondrial phylogenomics of the Bivalvia (Mollusca): searching for the origin and mitogenomic correlates of doubly uniparental inheritance of mtDNA. BMC Evol. Biol. 10, 1. https://doi.org/10.1186/1471-214810-50. Dowling, A., Clift, R., Grobert, N., Hutton, D., Oliver, R., O'neill, O., Pethica, J., Pidgeon, N., Porritt, J., Ryan, J., Seaton, A., Tendler, S., Welland, M., Whatmore, R., Enderby, J., Ruffles, P., Friend, R., Gilbert, N., McQuaid, J., Segal, A., 2004. Nanoscience and Nanotechnologies: Opportunities and Uncertainties. Lond. R. Soc. R. Acad. Eng. Rep. 46, 618. Dubois, M.K., Gilles, A., Hamilton, J.K., Rebers, P.A., Sith, F., 1956. Calorimetric method for determination of sugars and related substances. Anal. Chem. 28, 350–356. https://doi.org/10.1021/ac60111a017. EC, European Commission, 2011. Definition of a nanomaterial. http://ec.europa.eu/ environment/chemicals/nanotech/faq/definition_en.htm. Faggio, C., Pagano, M., Alampi, R., Vazzana, I., Felice, M.R., 2016. Cytotoxicity, haemolymphatic parameters, and oxidative stress following exposure to sub-lethal concentrations of quaternium-15 in Mytilus galloprovincialis. Aquat. Toxicol. 180, 258–265. https://doi.org/10.1016/j.marpolbul.2013.05.004. Fanslow, D.L., Nalepa, T.F., Johengen, T.H., 2001. Seasonal changes in the respiratory electron transport system (ETS) and respiration of the zebra mussel, Dreissena polymorpha in Saginaw Bay, Lake Huron. Hydrobiologia 448, 61–70. https://doi.org/10. 1023/A:1017582119098. Folch, J., Lees, M., Sloane Stanley, G.H., 1957. A simple method for the isolation and purification of total lipides from animal tissues. J. Biol. Chem. 226, 497–509. Freitas, R., Pires, A., Velez, C., Almeida, Â., Moreira, A., Wrona, F.J., Soares, A.M.V.M., Figueira, E., 2016a. Effects of seawater acidification on Diopatra neapolitana (Polychaete, Onuphidae): biochemical and regenerative capacity responses. Ecol. Indicat. 60, 152–161. https://doi.org/10.1016/j.ecolind.2015.06.032. Freitas, R., Salamanca, L., Velez, C., Wrona, F.J., Soares, A.M.V.M., Etelvina Figueira, E., 2016b. Multiple stressors in estuarine waters: Effects of arsenic and salinity on Ruditapes philippinarum. Sci. Total Environ. 541, 1106–1114. http://doi.org/10. 1016/j.scitotenv.2015.09.149. Freitas, R., De Marchi, L., Moreira, A., Pestana, J.L.T., Wrona, F.J., Figueira, E., Soares, A.M.V.M., 2017. Physiological and biochemical impacts induced by mercury pollution and seawater acidification in Hediste diversicolor. Sci. Total Environ. 595,

Al‐Shaeri, M., Ahmed, D., McCluskey, F., Turner, G., Paterson, L., Dyrynda, E.A., Hartl, M.G., 2013. Potentiating toxicological interaction of single‐walled carbon nanotubes with dissolved metals. Environ. Toxicol. Chem. 32 (12), 2701–2710. https://doi.org/ 10.1002/etc.2365. Andrade, M., Soares, A., Figueira, E., Freitas, R., 2018. Biochemical changes in mussels submitted to different time periods of air exposure. Environ. Sci. Pollut. Res. 25 (9), 8903–8913. https://doi.org/10.1007/s11356-017-1123-7. Andral, B., Stanisiere, J.Y., Sauzade, D., Damier, E., Thebault, H., Galgani, F., Boissery, P., 2004. Monitoring chemical contamination levels in the Mediterranean based on the use of mussel caging. Mar. Pollut. Bull. 49, 704–712. https://doi.org/10.1016/j. marpolbul.2004.05.008. Anestis, A., Lazou, A., Pörtner, H.-O., Michaelidis, B., 2007. Behavioral, metabolic, and molecular stress responses of marine bivalve Mytilus galloprovincialis during long-term acclimation at increasing ambient temperature. Aust. J. Pharm.: Regul. Integr. Comp. Physiol. 293, 911–921. https://doi.org/10.1152/ajpregu.00124.2007. Anisimova, A.A., Chaika, V.V., Kuznetsov, V.L., Golokhvast, K.S., 2015. Study of the influence of multiwalled carbon nanotubes (12–14 nm) on the main target tissues of the bivalve Modiolus modiolus. Nanotechnol. Russ. 10, 278–287. https://doi.org/10. 1134/S1995078015020020. Arndt, D.A., Moua, M., Chen, J., Klaper, R., 2013. Core structure and surface functionalization of carbon nanomaterials alter impacts to daphnid mortality, reproduction, and growth: acute assays do not predict chronic exposure impacts. Environ. Sci. Technol. 47 (16), 9444–9452. https://doi.org/10.1021/es4030595. Attig, H., Kamel, N., Sforzini, S., Dagnino, A., Jamel, J., Boussetta, H., Viarengo, A., Banni, M., 2014. Effects of thermal stress and nickel exposure on biomarkers responses in Mytilus galloprovincialis (Lam). Mar. Environ. Res. 94, 65–71. https://doi. org/10.1016/j.marenvres.2013.12.006. Azpeitia, K., Ferrer, L., Revilla, M., Pagaldai, J., Mendiola, D., 2016. Growth, biochemical profile, and fatty acid composition of mussel (Mytilus galloprovincialis Lmk.) cultured in the open ocean of the Bay of Biscay (northern Spain). Aquaculture 454, 95–108. http://doi.org/10.1016/j.aquaculture.2015.12.022. Balbi, T., Fabbri, R., Montagna, M., Camisassi, G., Canesi, L., 2017. Seasonal variability of different biomarkers in mussels (Mytilus galloprovincialis) farmed at different sites of the Gulf of La Spezia, Ligurian sea. Italy. Mar. Pollut. Bull. 116 (1–2), 348–356. https://doi.org/10.1016/j.marpolbul.2017.01.035. Banni, M., Hajer, A., Sforzini, S., Oliveri, C., Boussetta, H., Viarengo, A., 2014. Transcriptional expression levels and biochemical markers of oxidative stress in Mytilus galloprovincialis exposed to nickel and heat stress. Comp. Biochem. Physiol. C Toxicol. Pharmacol. 160, 23–29. https://doi.org/10.1016/j.cbpc.2013.11.005. Bayne, B.L., Newell, R.C., 1983. Physiological energetics of marine molluscs. In: Salevddin, A.S.M., Wilbur, K.M. (Eds.), The mollusca. Volume 4: Physiology, Part 1. Academic Press, New York, pp. 407–515. Beauchamp, C., Fridovich, I., 1971. Superoxide dismutase: improved assays and an assay applicable to acrylamide gels. Anal. Biochem. 44, 276–287. http://doi.org/10.1016/ 0003-2697(71)90370-8. Berridge, M.V., Herst, P.M., Tan, A.S., 2005. Tetrazolium dyes as tools in cell biology: new insights into their cellular reduction. Biotechnol. Annu. Rev. 11, 127–152. https:// doi.org/10.1016/S1387-2656(05)11004-7. Boukadida, K., Banni, M., Gourves, P.-Y., Cachot, J., 2016. High sensitivity of embryolarval stage of the Mediterranean mussel, Mytilus galloprovincialis to metal pollution in combination with temperature increase. Mar. Environ. Res. 122, 59–66. https:// doi.org/10.1016/j.marenvres.2016.09.007. Canesi, L., Ciacci, C., Betti, M., Fabbri, R., Canonico, B., Fantinati, A., Marcomini, A., Pojana, G., 2008. Immunotoxicity of carbon black nanoparticles to blue mussel hemocytes. Environ. Int. 34, 1114–1119. https://doi.org/10.1016/j.envint.2008.04. 002. Canesi, L., Fabbri, R., Gallo, G., Vallotto, D., Marcomini, A., Pojana, G., 2010a. Biomarkers in Mytilus galloprovincialis exposed to suspensions of selected nanoparticles (Nano carbon black, C60 fullerene, Nano-TiO2, Nano-SiO2). Aquat. Toxicol. 100, 168–177. https://doi.org/10.1016/j.aquatox.2010.04.009. Canesi, L., Ciacci, C., Vallotto, D., Marcomini, A., Pojana, G., 2010b. In vitro effects of suspensions of selected nanoparticles (C60 fullerene, TiO2, SiO2) on Mytilus hemocytes. Aquat. Toxicol. 96 (2), 151–158. https://doi.org/10.1016/j.aquatox.2009.10. 017. Carregosa, V., Figueira, E., Gil, A.M., Pereira, S., Pinto, J., Soares, A.M.V.M., Freitas, R., 2014a. Tolerance of Venerupis philippinarum to salinity: osmotic and metabolic aspects. Compar. Biochem. Physiol. A Mol. Integr. Physiol. 171, 36–43. https://doi.org/ 10.1016/j.cbpa.2014.02.009. Carregosa, V., Velez, C., Soares, A.M.V.M., Figueira, E., Freitas, R., 2014b. Physiological and biochemical responses of the polychaete Diopatra neapolitana to organic matter enrichment. Aquat. Toxicol. 155 (2014), 32–42. https://doi.org/10.1016/j.aquatox. 2014.05.029. Carregosa, V., Velez, C., Soares, A.M.V.M., Figueira, E., Freitas, R., 2014c. Physiological and biochemical responses of three Veneridae clams exposed to salinity changes. Comp. Biochem. Physiol. B Biochem. Mol. Biol. 177–178, 1–9. https://doi.org/10. 1016/j.cbpb.2014.08.001. Catsiki, V.-A., Florou, H., 2006. Study on the behavior of the heavy metals Cu, Cr, Ni, Zn, Fe, Mn and 137Cs in an estuarine ecosystem using Mytilus galloprovincialis as a bioindicator species: the case of Thermaikos gulf, Greece. J. Environ. Radioact. 86, 31–44. https://doi.org/10.1016/j.jenvrad.2005.07.005. Cattaruzza, M., Hecker, M., 2008. Protein carbonylation and decarboylation: a new twist to the complex response of vascular cells to oxidative stress. Circ. Res. 102, 273–274. Çelik, M.Y., Karayücel, S., Karayücel, I., Öztürk, R., Eyüboğlu, B., 2012. Meat yield,

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Marine Environmental Research 145 (2019) 11–21

M. Andrade, et al. 691–701. http://doi.org/10.1016/j.scitotenv.2017.04.005. García-Martín, E.E., McNeill, S., Serret, P., Leakey, R.J.G., 2014. Plankton metabolism and bacterial growth efficiency in offshore waters along a latitudinal transect between the UK and Svalbard. Deep Sea Res. Part 1 Oceanogr. Res. Pap. 92, 141–151. https://doi.org/10.1016/j.dsr.2014.06.004. Gestoso, I., Arenas, F., Olabarria, C., 2016. Ecological interactions modulate responses of two intertidal mussel species to changes in temperature and pH. J. Exp. Mar. Biol. Ecol. 474, 116–125. https://doi.org/10.1016/j.jembe.2015.10.006. Griffiths, C.L., Griffiths, R.J., 1987. Bivalvia. In: Pandian, T.J., Vernberg, F.J. (Eds.), Animal energetics. Volume 2: Bivalvia through Reptilia. Academic Press, New York, pp. 1–88. Habig, W.H., Pabst, M.J., Jakoby, W.B., 1974. Glutathione S-transferases. The first enzymatic step in mercapturic acid formation. J. Biol. Chem. 249, 7130–7139. Hotze, E.M., Phenrat, T., Lowry, G.V., 2010. Nanoparticle aggregation: challenges to understanding transport and reactivity in the environment. J. Environ. Qual. 39 (6), 1909. https://doi.org/10.2134/jeq2009.0462. Hu, M., Li, L., Sui, Y., Li, J., Wang, Y., Lu, W., Dupont, S., 2015. Effect of pH and temperature on antioxidant responses of the thick shell mussel Mytilus coruscus. Fish Shellfish Immunol. 46, 572–583. https://doi.org/10.1016/j.fsi.2015.07.025. Huang, X., Liu, Z., Xie, Z., Dupont, S., Huang, W., Wu, F., Kong, H., Liu, L., Sui, Y., Lin, D., Lu, W., Hu, M., Wang, Y., 2018. Oxidative stress induced by titanium dioxide nanoparticles increases under seawater acidification in the thick shell mussel Mytilus coruscus. Mar. Environ. Res. 137, 49–59. https://doi.org/10.1016/j.marenvres.2018. 02.029. IPCC, 2007. Climate change 2007: the physical science basis. In: Contribution of Work Group I to the Fourth Assessment Report of the Intergovernmental Panel on Climate. Cambridge University Press, Cambridge, UK. IPCC, 2014. Climate change 2014: Synthesis report. In: Contribution of Working Groups I, II and III to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change. IPMA, 2017. Portuguese Institute for Sea and Atmosphere (IPMA). https://www.ipma. pt/en/maritima/sat-sst/index-8days.jsp?area=zona2. Jackson, P., Jacobsen, N.R., Baun, A., Birkedal, R., Kühnel, D., Jensen, K.A., et al., 2013. Bioaccumulation and ecotoxicity of carbon nanotubes. Chem. Cent. J. 7 (1), 154–165. https://doi.org/10.1186/1752-153X-7-154. Jansen, J.M., Hummel, H., Bonga, S.W., 2009. The respiratory capacity of marine mussels (Mytilus galloprovincialis) in relation to the high temperature threshold. Comp. Biochem. Physiol. Mol. Integr. Physiol. 153, 399–440. https://doi.org/10.1016/j. cbpa.2009.03.013. Johansson, L.H., Borg, L.A.H., 1988. A spectrophotometric method for determination of catalase activity in small tissue samples. Anal. Biochem. 174, 331–336. http://doi. org/10.1016/0003-2697(88)90554-4. Kádár, E., Lowe, D.M., Solé, M., Fisher, A.S., Jha, A.N., Readman, J.W., Hutchinson, T.H., 2010. Uptake and biological responses to nano-Fe versus soluble FeCl3 in excised mussel gills. Anal. Bioanal. Chem. 396, 657–666. https://doi.org/10.1007/s00216009- 3191-0. Kamel, N., Attig, H., Dagnino, A., Boussetta, H., Banni, M., 2012. Increased temperatures affect oxidative stress markers and detoxification response to benzo[a]pyrene exposure in mussel Mytilus galloprovincialis. Arch. Environ. Contam. Toxicol. 63, 534–543. https://doi.org/10.1007/s00244-012-9790-3. Kamel, N., Burgeot, T., Banni, M., Chalghaf, M., Devin, S., Minier, C., Boussetta, H., 2014. Effects of increasing temperatures on biomarker responses and accumulation of hazardous substances in rope mussels (Mytilus galloprovincialis) from Bizerte lagoon. Environ. Sci. Pollut. Res. Int. 21 (9), 6108–6123. https://doi.org/10.1007/s11356014-2540-5. Kataoka, C., Nakahara, K., Shimizu, K., Kowase, S., Nagasaka, S., Ifuku, S., et al., 2016. Salinity-dependent toxicity of water-dispersible, single-walled carbon nanotubes to Japanese medaka embryos. J. Appl. Toxicol. 37 (4), 408–416. https://doi.org/10. 1002/jat.3373. Kefaloyianni, E., Gourgou, E., Ferle, V., Kotsakis, E., Gaitanaki, C., Beis, I., 2005. Acute thermal stress and various heavy metals induce tissue-specific pro- or anti-apoptotic avents via the p38-MAPK signal transduction pathway in Mytilus galloprovincialis (Lam.). J. Exp. Biol. 208, 4427–4436. https://doi.org/10.1242/jeb.01924. Keller, A.A., McFerran, S., Lazareva, A., Suh, S., 2013. Global life cycle releases of engineered nanomaterials. J. Nanoparticle Res. 15, 1692. https://doi.org/10.1007/ s11051-013-1692-4. King, F.D., Packard, T.T., 1975. Respiration and the activity of the respiratory electron transport system in marine zooplankton. Limnol. Oceanogr. 20, 849–854. http://doi. org/10.4319/lo.1975.20.5.0849. Köhler, A.R., Som, C., Helland, A., Gottschalk, F., 2008. Studying the potential release of carbon nanotubes throughout the application life cycle. J. Clean. Prod. 16, 927–937. https://doi.org/10.1016/j.jclepro.2007.04.007. Kristan, U., Kanduč, T., Osterc, A., Šlejkovec, Z., Ramšak, A., Stibilj, V., 2014. Assessment of pollution level using as a bioindicator species: The case of the Gulf of Trieste. Mar. Pollut. Bull. 89, 455–463. http://doi.org/10.1016/j.marpolbul.2014.09.046. Letendre, J., Dupont-Rouzeyrol, M., Hanquet, A., Durand, F., Budzinski, H., Chan, P., Vaudry, D., Rocher, B., 2011. Impact of toxicant exposure on the proteomic response to intertidal condition in Mytilus edulis. Comp. Biochem. Physiol. Genom. Proteonom. 6 (4), 357–369. https://doi.org/10.1016/j.cbd.2011.08.002. Liné, C., Larue, C., Flahaut, E., 2017. Carbon nanotubes: Impacts and behaviour in the terrestrial ecosystem-A review. Carbon 123, 767–785. https://doi.org/10.1016/j. carbon.2017.07.089. Liu, C., Liang, C., Lin, K., Jang, C., Wang, S., Huang, Y., Hsueh, Y., 2007. Bioaccumulation of arsenic compounds in aquacultural clams (Meretrix lusoria) and assessment of potential carcinogenic risks to human health by ingestion. Chemosphere 69, 128–134. https://doi.org/10.1016/j.chemosphere.2007.04.038.

Lucas, A., Beninger, P.G., 1985. The use of physiological condition indices in marine bivalve aquaculture. Aquaculture 44, 187–200. https://doi.org/10.1016/00448486(85)90243-1. Lushchak, V.I., 2011. Environmentally induced oxidative stress in aquatic animals. Aquat. Toxicol. 101, 13–30. https://doi.org/10.1016/j.aquatox.2010.10.006. Marisa, I., Matozzo, V., Munari, M., Binelli, A., Parolini, M., Martucci, A., 2016. In vivo exposure of the marine clam Ruditapes philippinarum to zinc oxide nanoparticles: responses in gills, digestive gland and haemolymph. Environ. Sci. Pollut. Res. 23, 15275. https://doi.org/10.1007/s11356-016-6690-5. Martínez-Álvarez, R.M., Morales, A.E., Sanz, A., 2005. Antioxidant defenses in fish: biotic and abiotic factors. Rev. Fish Biol. Fish. 15, 75–88. http://doi.org/10.1007/s11160005-7846-4. Matozzo, V., Binelli, A., Parolini, M., Previato, M., Masiero, L., Finos, L., Bressan, M., Marin, M.G., 2012. Biomarker responses in the clam Ruditapes philippinarum and contamination levels in sediments from seaward and landward sites in the lagoon of Venice. Ecol. Indicat. 19, 191–205. http://doi.org/10.1016/j.ecolind.2011.06.020. Mesquita, C.S., Oliveira, R., Bento, F., Geraldo, D., Rodrigues, J.V., Marcos, J.C., 2014. Simplified 2,4-dinitrophenylhydrazine spectrophotometric assay for quantification of carbonyls in oxidized proteins. Anal. Biochem. 458, 69–71. https://doi.org/10.1016/ j.ab.2014.04.034. Miller, M.A., Bankier, C., Al-Shaeri, M.A.M., Hartl, M.G.J., 2015. Neutral red cytotoxicity assays for assessing in vivo carbon nanotube ecotoxicity in mussels—Comparing microscope and microplate methods. Mar. Pollut. Bull. 101 (2), 903–907. https://doi. org/10.1016/j.marpolbul.2015.10.072. Mitrano, D.M., Motellier, S., Clavaguera, S., Nowack, B., 2015. Review of nanomaterial aging and transformations through the life cycle of nano-enhanced products. Environ. Int. 77, 132–147. https://doi.org/10.1016/j.envint.2015.01.013. Moore, M.N., 2006. Do nanoparticles present ecotoxicological risks for the health of the aquatic environment? Environ. Int. 32, 967–976. https://doi.org/10.1016/j.envint. 2006.06.014. Moore, M.N., Readman, J.A.J., Readman, J.W., Lowe, D.M., Frickers, P.E., Beesley, A., 2009. Lysosomal cytotoxicity of carbon nanoparticles in cells of the molluscan immune system: an in vitro study. Nanotoxicology 3, 40–45. https://doi.org/10.1080/ 17435390802593057. Moschino, V., Nesto, N., Barison, S., Agresti, F., Colla, L., Fedele, L., Da Ros, L., 2014. A preliminary investigation on nanohorn toxicity in marine mussels and polychaetes. Sci. Total Environ. 468, 111–119. https://doi.org/10.1016/j.scitotenv.2013.08.020. Muller, N., Nowack, B., 2008. Exposure modeling of engineered nanoparticles in the environment. Environ. Sci. Technol. 41, 4447–4453. https://doi.org/10.1021/ es7029637. Mwangi, J.N., Wang, N., Ingersoll, C.G., Hardesty, D.K., Brunson, E.L., Li, H., Deng, B., 2012. Toxicity of carbon nanotubes to freshwater aquatic invertebrates. Environ. Toxicol. Chem. 31 (8), 1823–1830. Nardi, A., Mincarelli, L.F., Benedetti, M., Fattorini, D., d'Errico, G., Regoli, F., 2017. Indirect effects of climate changes on cadmium bioavailability and biological effects in the Mediterranean mussel Mytilus galloprovincialis. Chemosphere 169, 493–502. https://doi.org/10.1016/j.chemosphere.2016.11.093. Nilin, J., Pestana, J.L.T., Ferreira, N.G., Loureiro, S., Costa-Lotufo, L.V., Soares, A.M.V.M., 2012. Physiological responses of the European cockle Cerastoderma edule (Bivalvia: Cardidae) as indicators of coastal lagoon pollution. Sci. Total Environ. 435–436, 44–52. http://doi.org/10.1016/j.scitotenv.2012.06.107. Ohkawa, H., Ohishi, N., Yagi, K., 1979. Assay for lipid peroxides in animal tissues by thiobarbituricacid reaction. Anal. Biochem. 95, 351–358. http://doi.org/10.1016/ 0003-2697(79)90738-3. Oliveira, P., Almeida, Â., Calisto, V., Esteves, V.I., Schneider, R.J., Wrona, F.J., Soares, A.M.V.M., Figueira, E., Freitas, R., 2017. Physiological and biochemical alterations induced in the mussel Mytilus galloprovincialis after short and long-term exposure to carbamazepine. Water Res. 117, 102–114. https://doi.org/10.1016/j.watres.2017. 03.052. Orban, E., Di Lena, G., Nevigato, T., Casini, I., Marzetti, A., Caproni, R., 2002. Seasonal changes in meat content, condition index and chemical composition of mussels (Mytilus galloprovincialis) cultured in two different Italian sites. Food Chem. 77, 57–65. https://doi.org/10.1016/S0308-8146(01)00322-3. Paglia, D.E., Valentine, W.N., 1967. Studies on the quantitative and qualitative characterization of erythrocyte glutathione peroxidase. J. Lab. Clin. Med. 70 (1), 158–169. https://doi.org/10.1002/etc.710. Paital, B., Chainy, G.B.N., 2013. Seasonal variability of antioxidant biomarkers in mud crabs (Scylla serrata). Ecotoxicol. Environ. Saf. 87, 33–41. https://doi.org/10.1016/j. ecoenv.2012.10.006. Peng, X., Jia, J., Gong, X., Luan, Z., Fan, B., 2009. Aqueous stability of oxidized carbon nanotubes and the precipitation by salts. J. Hazard Mater. 165, 1239–1242. https:// doi.org/10.1016/j.jhazmat.2008.10.049. Petersen, E.J., Zhang, L., Mattison, N.T., O'Carroll, D.M., Whelton, A.J., Uddin, N., Nguyen, T., Huang, Q., Henry, T.B., Holbrook, R.D., Chen, K.L., 2011. Potential release pathways, environmental fate, and ecological risks of carbon nanotubes. Environ. Sci. Technol. 45, 9837–9856. https://doi.org/10.1021/es201579y. Pörtner, H.-O., 2010. Oxygen- and capacity-limitation of thermal tolerance: a matrix for integrating climate-related stressor effects in marine ecosystems. J. Exp. Biol. 213, 881–893. https://doi.org/10.1242/jeb.037523. Pörtner, H.-O., Knust, R., 2007. Climate change affects marine fishes through the oxygen limitation of thermal tolerance. Science 315, 95–97. https://doi.org/10.1126/ science.1135471. Pörtner, H.-O., Langenbuch, M., Michaelidis, B., 2005. Synergistic effects of temperature extremes, hypoxia, and increases in CO2 on marine animals: from earth history to global change. J. Geophys. Res. 110, 2156–2202. https://doi.org/10.1029/ 2004JC002561.

20

Marine Environmental Research 145 (2019) 11–21

M. Andrade, et al. Potočnik, J., 2011. Commission recommendation of 18 October 2011 on the definition of nanomaterial (2011/696/EU). Official Journal of the European Union L275, 38–40. https://doi.org/10.2777/13162. Regoli, F., Giuliani, M.E., 2014. Oxidative pathways of chemical toxicity and oxidative stress biomarkers in marine organisms. Mar. Environ. Res. 93, 106–117. https://doi. org/10.1016/j.marenvres.2013.07.006. Relexans, J.C., Meybeck, M., Billen, G., Brugeaille, M., Etcheber, H., Somville, M., 1988. Algal and microbial processes involved in particulate organic matter dynamics in the Loire estuary. Estuar. Coast Shelf Sci. 27, 625–644. http://doi.org/10.1016/ 02727714(88)90072-8. Resgalla Jr., C., Brasil, E.S., Salomao, L.C., 2007. The effect of temperature and salinity on the physiological rates of the mussel Perna perna (Linnaeus 1758). Braz. Arch. Biol. Technol. 50 (3), 543–556. http://doi.org/10.1590/S1516-89132007000300019. Resgalla, C., Radetski, C.M., Schettini, C.A.F., 2010. Physiological energetics of the brown mussel Perna perna (L.) transplanted in the Itajaì-Açu river mouth, Southern Brazil. Ecotoxicology 19, 383–390. http://doi.org/10.1007/s10646-009-0422-2. Rocha, T.L., Gomes, T., Sousa, V.S., Mestre, N.C., Bebianno, M.J., 2015. Ecotoxicological impact of engineered nanomaterials in bivalve molluscs: An overview. Mar. Environ. Res. 111, 74–88. https://doi.org/10.1016/j.marenvres.2015.06.013. Sanchez, V.C., Jachak, A., Hurt, R.H., Kane, A.B., 2012. Biological interactions of graphene-family nanomaterials–an interdisciplinary review. Chem. Res. Toxicol. 15–34. https://doi.org/10.1021/tx200339h.Biological. Santos, A.L., Mendes, C., Gomes, N.C.M., Henriques, I., Correia, A., Almeida, A., Cunha, Â., 2009. Short-term variability of abundance, diversity and activity of estuarine bacterioneuston and bacterioplankton. J. Plankton Res. 31, 1545–1555. https://doi. org/10.1093/plankt/fbp083. Scott-Fordsmand, J.J., Weeks, J.M., 2000. Biomarkers in earthworms. Rev. Environ. Contam. Toxicol. 165, 117–159. http://doi.org/10.1007/978-1-4612-1172-3_3. Silva, A.Z., Zanette, J., Ferreira, J.F., Guzenski, J., Marques, M.R.F., Bainy, A.C.D., 2005. Effects of salinity on biomarker responses in Crassostrea rhizophorae (Mollusca, Bivalvia) exposed to diesel oil. Ecotoxicol. Environ. Saf. 62, 376–382. http://doi.org/ 10.1016/j.ecoenv.2004.12.008. Solarskaciuk, K., Gajewska, A., Glińska, S., Michlewska, S., Balcerzak, Ł., Jamrozik, A., Skolimowski, J., Burda, K., Bartosz, G., 2014. Effect of functionalized and nonfunctionalized nanodiamond on the morphology and activities of antioxidant enzymes of lung epithelial cells (A549). Chem. Biol. Interact. 222, 135–147. https://doi. org/10.1016/j.cbi.2014.10.003. Sultan, C.S., Saackel, A., Stank, A., Fleming, T., Fedorova, M., Hoffmann, R., Wade, R.C., Hecker, M., Wagner, A.H., 2018. Impact of carbonylation on glutathione peroxidase-1 activity in human hyperglycemic endothelial cells. Redox Biol 16, 113–122. https:// doi.org/10.1016/j.redox.2018.02.018. Sun, Y., Fu, K., Lin, Y.I., 2002. Functionalized carbon nanotubes: properties and applications. Acc. Chem. Res. 35 (12), 1096–1104. https://doi.org/10.1021/ar010160v. Sun, T.Y., Bornhöft, N.A., Hungerbühler, K., Nowack, B., 2016. Dynamic probabilistic modeling of environmental emissions of engineered nanomaterials. Environ. Sci. Technol. 50, 4701–4711.

Suzuki, Y.J., Carini, M., Butterfield, D.A., 2010. Protein Carbonylation. Antioxid. Redox Signal 12 (3), 323–325. https://doi.org/10.1089/ars.2009.2887. Velez, C., Figueira, E., Soares, A.M.V.M., Freitas, R., 2015. Spatial distribution and bioaccumulation patterns in three clam populations from a low contaminated ecosystem. Estuar. Coast Shelf Sci. 155, 114–125. https://doi.org/10.1016/j.ecss.2015. 01.004. Velez, C., Figueira, E., Soares, A.M.V.M., Freitas, R., 2016. Combined effects of seawater acidification and salinity changes in Ruditapes philippinarum. Aquat. Toxicol. 176, 141–150. http://doi.org/10.1016/j.aquatox.2016.04.016. Velez, C., Figueira, E., Soares, A.M.V.M., Freitas, R., 2017. Effects of seawater temperature increase on economically relevant native and introduced clam species. Mar. Environ. Res. 123, 62–70. https://doi.org/10.1016/j.marenvres.2016.11.010. Verlecar, X.N., Jena, K.B., Chainy, G.B.N., 2007. Biochemical markers of oxidative stress in Perna viridis exposed to mercury and temperature. Chem. Biol. Interact. 167 (3), 219–226. https://doi.org/10.1016/j.cbi.2007.01.018. Viarengo, A., Dondero, F., Pampanin, D.M., Fabbri, R., Poggi, E., Malizia, M., Bolognesi, C., Perrone, E., Gollo, E., Cossa, G.P., 2007. A biomonitoring study assessing the residual biological effects of pollution caused by the HAVEN wreck on marine organisms in the Ligurian sea (Italy). Arch. Environ. Contam. Toxicol. 53, 607–616. https://doi.org/10.1007/s00244-005-0209-2. Vlasova, I.I., Kapralov, A.A., Michael, Z.P., Burkert, S.C., Shurin, M.R., Star, A., Shvedova, A.A., Kagan, V.E., 2016. Enzymatic oxidative biodegradation of nanoparticles: mechanisms, significance and applications. Toxicol. Appl. Pharmacol. 299, 58–69. https://doi.org/10.1016/j.taap.2016.01.002. Wang, W.X., Fisher, N.S., Luoma, S.N., 1996. Kinetic determinations of trace element bioaccumulation in the mussel Mytilus edulis. Mar. Ecol. Prog. Ser. 140, 91–113. https://doi.org/10.3354/meps140091. Wang, Y., Li, L., Hu, M., Lu, W., 2015. Physiological energetics of the thick shellmussel Mytilus coruscus exposed to seawater acidification and thermal stress. Sci. Total Environ. 514, 261–272. http://doi.org/10.1016/j.scitotenv.2015.01.092. Ward, J.E., Kach, D.J., 2009. Marine aggregates facilitate ingestion of nanoparticles by suspension-feeding bivalves. Mar. Environ. Res. 68 (3), 137–142. https://doi.org/10. 1016/j.marenvres.2009.05.002. Wu, Q., Yin, L., Li, X., Tang, M., Zhang, T., Wang, D., 2013. Contributions of altered permeability of intestinal barrier and defecation behavior to toxicity formation from graphene oxide in nematode Caenorhabditis elegans. Nanoscale 5, 9934–9943. https:// doi.org/10.1039/c3nr02084c. Yang, J.L., Li, Y.F., Guo, X.P., Liang, X., Xu, Y.F., Ding, D.W., Bao, W.-Y., Dobretsov, S., 2016. The effect of carbon nanotubes and titanium dioxide incorporated in PDMS on biofilm community composition and subsequent mussel plantigrade settlement. Biofouling 32 (7), 763–777. Zhang, X., Zhou, Q., Zou, W., Hu, Xiangang, H., 2017. Molecular Mechanisms of Developmental Toxicity Induced by Graphene Oxide at Predicted Environmental Concentrations. Environ. Sci. Technol. 51 (14), 7861–7871. https://doi.org/10. 1021/acs.est.7b01922.

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